Wu et al. AMB Expr (2018) 8:27 https://doi.org/10.1186/s13568-018-0560-0

ORIGINAL ARTICLE

Biodegradation of decabromodiphenyl ether (BDE 209) by a newly isolated bacterium from an e-waste recycling areaZhineng Wu1, Miaomiao Xie1, Yao Li1, Guanghai Gao1, Mark Bartlam2,3 and Yingying Wang1*

Abstract

Polybrominated diphenyl ethers (PBDEs) have become widespread environmental pollutants all over the world. A newly isolated bacterium from an e-waste recycling area, Stenotrophomonas sp. strain WZN-1, can degrade decabro-modiphenyl ether (BDE 209) effectively under aerobic conditions. Orthogonal test results showed that the optimum conditions for BDE 209 biodegradation were pH 5, 25 °C, 0.5% salinity, 150 mL minimal salt medium volume. Under the optimized condition, strain WZN-1 could degrade 55.15% of 65 μg/L BDE 209 under aerobic condition within 30 day incubation. Moreover, BDE 209 degradation kinetics was fitted to a first-order kinetics model. The biodegradation mechanism of BDE 209 by strain WZN-1 were supposed to be three possible metabolic pathways: debromination, hydroxylation, and ring opening processes. Four BDE 209 degradation genes, including one hydrolase, one dioxy-genase and two dehalogenases, were identified based on the complete genome sequencing of strain WZN-1. The real-time qPCR demonstrated that the expression level of four identified genes were significantly induced by BDE 209, and they played an important role in the degradation process. This study is the first to demonstrate that the newly isolated Stenotrophomonas strain has an efficient BDE 209 degradation ability and would provide new insights for the microbial degradation of PBDEs.

Keywords: Stenotrophomonas, Biodegradation, BDE 209, Genome sequencing, Real-time qPCR

© The Author(s) 2018. This article is distributed under the terms of the Creative Commons Attribution 4.0 International License (http://creativecommons.org/licenses/by/4.0/), which permits unrestricted use, distribution, and reproduction in any medium, provided you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons license, and indicate if changes were made.

IntroductionPBDEs are extensively used in various types of electronic equipment, furniture, plastics, textiles, etc. due to their excellent flame retardant performance (Leonel et  al. 2014). Since PBDEs can be easily incorporated into dif-ferent materials and products without making any cova-lent bonds, they could transport and migrate to different environment through volatilization and effusion etc. (Kim et al. 2007). Therefore, PBDEs have become wide-spread environmental pollutants and are ubiquitous in the environment (Birgul et  al. 2012). Due to their per-sistency, long-range environmental transport potential and bioaccumulation ability, four PBDE congeners (i.e.

tetrabromodiphenyl ether, pentabromodiphenyl ether, hexabromodiphenyl ether and heptabromodiphenyl ether) have been listed as new persistent organic pol-lutants (POPs) in 2009 (Stockholm Convention 2009). Previous studies demonstrated that PBDEs have a wide range of toxicities, including cytotoxicity (An et al. 2011), developmental neurotoxicity (Costa et  al. 2014), geno-toxicity (Ji et  al. 2011), disruption of thyroid functions (Xu et  al. 2014), reproductive toxicity (Muirhead et  al. 2006) and carcinogenicity (Zhao et  al. 2015). Penta-BDEs are reported to be the most toxic compounds among the PBDEs congeners. Deca-BDEs received the least concern among PBDEs congeners on account of its relatively lower toxicity (Gandhi et  al. 2011). How-ever, the higher-brominated PBDEs congeners (hepta-BDEs to deca-BDEs) could be degraded/transformed to lower-brominated PBDEs congeners which seems to be more toxic in various environments such as soil, aquatic ecosystems (Huang et  al. 2014). Therefore, PBDEs had

Open Access

*Correspondence: wangyy@nankai.edu.cn 1 Key Laboratory of Pollution Processes and Environmental Criteria (Ministry of Education), Tianjin Key Laboratory of Environmental Remediation and Pollution Control, College of Environmental Science and Engineering, Nankai University, Tianjin 300350, ChinaFull list of author information is available at the end of the article

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posed a great threat to both human health and the global environment.

PBDEs are lipophilic, and have very low solubility in water (ng/L to μg/L) (He et  al. 2006). Hence, accumu-lation of PBDEs in the environment, as well as in fatty tissues of humans and other animals are very likely to happen. Therefore, increased number of research have focused on the distribution and level of PBDEs in dif-ferent environments (Kim et  al. 2014), biota (Vorkamp et al. 2011) and in humans (Abdallah and Harrad 2014) worldwide. In contrast, there are very few reports on the degradation of PBDEs, especially on the biodegrada-tion of PBDE compounds. Given their worldwide distri-bution, it is important to develop effective degradation methods to remove PBDEs from environment, especially the most abundant congener BDE 209 (Ma et  al. 2012; Mcgrath et  al. 2017). The reported microbial degrada-tion of PBDEs have focused on the low PBDE congeners. For instance, a previous study showed Sphingomonas sp. SS3 can transform and grow on 4-bromodiphenyl ether (Schmidt et al. 1993). Robrock et al. reported the bacte-rial transformation of mono- to hexa-BDEs at μg/L lev-els (Robrock et  al. 2011). More recently, Pseudomonas stutzeri, Pseudomonas putida sp. strain TZ-1 and Pseu-domonas sp. were used to biodegrade different low PBDE congeners (Xin et al. 2014; Yang et al. 2015; Zhang et al. 2013). To the best of our knowledge, little information on microbial degradation of BDE 209 is available under aerobic condition.

In the present study, we investigated the BDE 209 biodegradation by a newly isolated bacterium from an e-waste recycling area in Tianjin, China. This bacterial strain can utilize BDE 209 as its carbon source effec-tively under aerobic conditions. Orthogonal tests were designed to acquire optimal condition for biodegradation of BDE 209. A first-order kinetics model was used to fit the BDE 209 degradation kinetics. Moreover, the inter-mediate products were detected by GC–MS and UPLC–MS. Furthermore, key BDE 209 degradation genes were identified through the complete genome sequencing analysis as well as real-time qPCR tests.

Materials and methodsChemicalsBDE 209 standard solution (concentration 50  µg/mL) in isooctane: toluene (9:1, v: v) and standard solutions containing 14 PBDE congeners (concentration 5 µg/mL, decabromodiphenyl ether was 25  µg/mL) in isooctane were purchased from AccuStandard (New Haven, CT, US). BDE 209 (99% purity) was bought from J&K Sci-entific Ltd and dissolved in tetrahydrofuran (THF) as stock solutions for bacteria isolation and biodegradation

experiments. Isooctane was HPLC grade, and the other chemicals and solvents used in the experiments were analytical grade.

Bacterial isolationSoil samples (0–15  cm) were collected in November 2014 from Ziya e-waste recycling area in Tianjin, north China (38°52′33.1″N, 116°49′48.7″E). Bacterium capa-ble of using BDE 209 as the carbon source was enriched and isolated. The minimal salt medium (MSM) con-tained basal mineral medium (BMM) (Na2HPO4·12H2O 25.6 g/L, KH2PO4 3 g/L, (NH4)2SO4 1.77 g/L), and trace mineral solution (1% v/v) containing HCl 20  mL/L, CaCO3 8  g/L, FeCl3·6H2O 7.74  g/L, MnCl2·4H2O 1.15 g/L, CuSO4·5H2O 0.146 g/L, CoCl2·6H2O 0.13 g/L, ZnO 0.4 g/L, H3BO3 0.124 g/L, EDTA Na4·2H2O 79.2 g/L, MgCl2·6H2O 13.42  g/L, Na2MoO4·2H2O 1.04  g/L. The pH of MSM was 7.4.

5 g soil sample was suspended in 50 mL sterilized saline solution (0.85%, w: v), and vortexed for 5 min, followed by 10  min standing at room temperature. The superna-tant was passed through 0.45 μm sterile filters, and then 1 mL filtrate was transferred to sterilized MSM contain-ing 10  mg/L BDE 209. The Erlenmeyer flasks (250  mL) were incubated 30 ± 1 °C, 150 rpm for 7 days. After three rounds of the supernatant transferring, which means transfer MSM with lower BDE 209 concentration to fresh MSM with higher BDE 209 concentration, the BDE 209 concentration finally reached 30 mg/L. Then, the enrich-ment culture was streaked in solid LB (Luria–Bertani) medium plates. A single colony on plates were picked and inoculated into LB medium for further identification and BDE 209 degradation. The isolated bacterium was designated as strain WZN-1.

Bacterial identificationThe isolated bacterium was identified based on 16S rRNA sequence analysis and morphological properties. The genomic DNA of strain WZN-1 was extracted by the UltraClean® Microbial DNA Kit (MoBio, USA). The 16S rRNA gene sequence was amplified by PCR using the universal primers (Lane 1991). The PCR products were sent to the AuGCT DNA-SYN Biotechnology Co., Ltd. (Beijing, China) for sequencing. The sequence was aligned and compared with those in the GenBank data-base, using ClustalW with default parameter values. A phylogenetic tree was constructed by the neighbor-join-ing method using the program MEGA 6.0 software. The morphological characteristics were examined by scan-ning electron microscopy (SEM) using an S-3500 N SEM (Hitachi, Japan).

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Optimization of BDE 209 degradation conditionsStrain WZN-1 was grown in 250  mL Erlenmeyer flask containing 100  mL of LB medium and incubated at 35 ± 1 °C for 36 h until its late-exponential phase. Then harvested by centrifugation at 7000  rpm for 5  min, and washed with 0.85% sterilized saline solution for three times. The initial bacterial concentration was 2  ×  104 cells/mL. Orthogonal test of five factors with five levels was used to identify the optimal BDE 209 degradation condition (Xu et al. 2011). The selected five factors were pH, temperature, salinity, MSM volume and initial BDE 209 concentration, respectively. The different levels of the five selected factors were shown in Additional file 1: Table S1. A total of 25 experimental combinations were carried out (Additional file  1: Table S2). All the experi-ments were set in triplicate. The BDE 209 concentration was measured after 7 days incubation. The optimal BDE 209 degradation conditions were determined by analyz-ing the orthogonal experiment results.

Biodegradation of BDE 209 by strain WZN‑1 under the optimal conditionBatch biodegradation experiments were performed in 250 mL serum bottles under the optimized BDE 209 deg-radation condition, temperature 25 ±  1  °C, 0.5% salin-ity, 150  mL MSM volume. BDE 209 concentration was 65 μg/L. The pH was adjusted to 7 by HCl, because strain WZN-1 showed better growth at pH 7 than pH 5 accord-ing to the orthogonal test. The initial bacterial concen-tration was 6  ×  104 cells/mL. Un-inoculated control bottles were kept to account for the abiotic loss of BDE 209. Glassware used in this experiments were carbon free (Vital et  al. 2012). All experiments were carried out in triplicate. Samples were taken at regular time intervals. Bacterial growth was monitored by FCM (flow cytom-etry) and BDE 209 concentration was determined by GC–MS. The FCM analysis was carried out on a CyFlow Space instrument (Partec, Hamburg, Germany). Detailed procedure was as previously described (Hammes et  al. 2008; Prest et  al. 2013; Wang et  al. 2009). The specific instrumental gain settings were as follows: green fluo-rescence  =  250, red fluorescence  =  700, SSC  =  250, FSC =  700. The same setting was used for all samples tested in this study.

The first-order kinetics model was used to analyze BDE 209 biodegradation data by strain WZN-1 (Eq. 1) (Peng et al. 2015).

where C0 is the initial BDE 209 concentrations, Ct refer to BDE209 concentration at time t; k and t are the degrada-tion rate constant and degradation time, respectively.

(1)Ct

C0

= e−kt

The half-lives (T1/2) of BDE 209 biodegradation were determined by the algorithm (Eq. 2) (Peng et al. 2015).

Intermediate products extraction and analysisSamples containing PBDEs were extracted by liquid–liq-uid extraction (Lee and He 2010). The extraction effi-ciencies of different PBDEs congeners, measured as the amount of PBDEs mixture recovered compared to the amount of PBDEs added to the control samples, with an extraction variation range of 83.09–103.21%. The BDE 209 concentration and the 13 lower PBDEs congeners were identified and quantified by GC–MS (7890A-5975C, Agilent Technology, USA) equipped with Rtx-1614 GC column (15 m × 0.25 mm × 0.10 μm). Automatic split-less injection was used with 1 μL injection volume at 280 °C. The carrier gas was helium (flow rate 2 mL/min). Pressure was kept constant at 21.77 psi. The oven tem-perature was started at 110  °C for 3  min, and then was heated to 200 °C at a rate of 25 °C/min, 15 °C/min up to 280 °C, and finally 20 °C/min increased to 305 °C held for 5 min. The total program time was 18.183 min. The ion source temperature was set at 150  °C, and the auxiliary temperature was 300  °C. Negative chemical ionization (NCI) mode was used in mass spectrometry, reagent gas was methane, and the ionization energy was 70 eV. The detection was selected ion monitoring in the NCI mode. The m/z 486.7 and 488.7 were monitored ions for BDE 209, and ions m/z 79 and 81 were selected for the 13 lower PBDEs congeners except BDE 209.

1 mL of the extracted top organic layer (method is the same with BDE 209 extraction) was transferred to dry test tubes, dried under nitrogen, then dissolved in 1 mL of methanol, and filtered through a 0.22  μm membrane filter. The UPLC–ESI–MS method was developed for the determination of OH-PBDEs (Wang et al. 2011). UPLC–MS (Waters ACQUITY UPLC H-CLASS, Waters, Mil-ford, MA, USA) equipped with an ESI source (Waters, Milford, MA, USA) was used to identify intermediate products. The parameters were as follows: capillary volt-age, 10 kV; source temperature, 160 °C; desolvation tem-perature, 400 °C. The mass spectrometer was operated in negative electrospray ionization (ESI-) mode using meth-anol–water (v: v, 90/10) as the mobile phase with a flow rate of 2 mL/min.

Complete genome sequencing and degradation genes identificationThe genomic DNA of strain WZN-1 was extracted as described above in “Bacterial identification” section. The complete genome sequencing of WZN-1 was finished using Illumina Hiseq  4000 and PacBio RSII sequenc-ing platform. The detailed sequencing information were

(2)T1/2 = ln 2/k

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described in a previous study (Wu et al. 2017). The spe-cific degradation genes were identified and searched from the annotation of all the genes in WZN-1 use key words (such as dehalo, ring-opening) for further real-time qPCR analysis.

Real‑time qPCRStrain WZN-1 was incubated in MSM medium supple-mented with 65 μg/L BDE 209 under the optimized BDE 209 degradation conditions (same with “Biodegradation of BDE 209 by strain WZN-1 under the optimal condi-tion batch” section). Strain WZN-1 was grown in MSM medium without BDE 209 was set as the control. The initial bacterial concentration was 1.36  ×  106 cells/mL so that the biomass can meet the need for RNA extrac-tion. Samples were collected in 5, 10, and 20  days from the bottles for the FCM analysis and real-time qPCR analysis. The total RNA of the 10 mL liquid samples (107 cells/mL) were extracted using the Total RNA Purifica-tion Kit (GeneMark, Taichung, Taiwan) as described by the manufacturer. Genomic DNA was eliminated by RNase-free DNase I treatment during the extraction pro-cedure. The SAMreal™ One-Step RT qPCR Kit (SYBR Green with low ROX) (GeneMark, Taichung, Taiwan) was used for the real time qPCR experiments in 96-well plates performed on the LightCycler® 96 real-time PCR system (Roche, Switzerland). The primers for real-time PCR were designed with the Primer Premier 5.0 soft-ware and synthesized by Sangon Biotech Co (Shanghai, China) (Table 1). The 16S rRNA gene was used as a refer-ence gene to normalize gene expression in strain WZN-1 (Wang and Shao 2012). Each real-time qPCR mixture (20 µL) contained 1 μL RNA, 1 μL enzyme mix, 1 0 μL 2× RT qPCR reaction (low ROX), 0.8 μL primers and 7.2 μL Rnase-free water. For each qPCR experiment, con-trol reactions without reverse transcriptase were con-ducted to verify the absence of genomic DNA. Cycling parameters were designed as follows: reverse transcrip-tion at 50  °C for 3 min, initial denaturation at 95  °C for 10 min, followed by 40 cycles of 95 °C for 5 s and 60 °C for 30  s. The fluorescence values were measured during

each annealing step. The melting curve was used to verify the specificity of the real-time qPCR reaction. The rela-tive fold change in mRNA quantity was calculated for the gene of interest in each sample using the 2−ΔΔCt method (Livak and Schmittgen 2001). All experiments were car-ried out in triplicate.

Strain accession numberThe 16S rRNA sequence of strain WZN-1 was depos-ited in NCBI GeneBank with the accession number of KY000522. The complete genome sequence of strain WZN-1 have been deposited in NCBI with a GenBank accession number of CP021768.

ResultsIdentification of strain WZN‑1A pure strain that can utilize BDE 209 as its carbon source was isolated from the mixed culture after 1 month of enrichment. This strain was designated as WZN-1. After incubation at 37 °C for 2 days, the colonies of strain WZN-1 on LB medium agar plates were approximately 1–2  mm in diameter. Cells of strain WZN-1 were rod-shaped, and measured 0.25–0.7 μm by 0.3–2 μm (Addi-tional file  1: Fig. S1a). This strain was a Gram-negative bacterium. The physiological and biochemical charac-teristics of strain WZN-1 are shown in Additional file 1: Table S3. According to a BLAST search against the NCBI GenBank, the 16S rRNA sequence of strain WZN-1 has a high sequence similarity with Stenotrophomonas (Addi-tional file 1: Fig. S2). The taxonomic position showed that the strain WZN-1 was a member of the Stenotropho-monas subgroup in the class γ-proteobacteria. The col-lection number of Stenotrophomonas sp. strain WZN-1 was CGMCC No. 12918 in China General Microbiologi-cal Culture Collection Center. Furthermore, this strain exhibited growth properties as a facultative oligotroph. The strain exhibited good growth at a wide range of car-bon concentrations (13–13,000  mg C/L) (Additional file  1: Fig. S3). Compared with Escherichia coli BL21, strain WZN-1 demonstrated obvious growth advan-tages under oligotrophic condition. According to the

Table 1 Genes and corresponding primers for the real-time qPCR

Gene name Gene ID Gene function Forward (5′ to 3′) Reverse (5′ to 3′) References

16S rRNA – Reference gene CGGTGAATACGTTCYCGG GGWTACCTTGTTACGACTT Wang et al. (2015)

Hydrolase CCR98_00905 Alpha/beta hydrolase AGCTATTACTGGCGCACGTT TGTACTCGTAGCGGCTGTCA Wu et al. (2017)

Dioxygenase CCR98_02495 Aromatic ring-opening dioxygenase

AAGGCCGAGCAGGATTATCT GACAGCGTCATGCTCTTCAC Wu et al. (2017)

Dehalogenase gene 1 CCR98_02005 Haloacid dehalogenase ATCTGTTCGCCTCGCTGAT ATAGACCGAAATGCCAGCAC Wu et al. (2017)

Dehalogenase gene 2 CCR98_19135 Haloacid dehalogenase CAGCATCACCCACAACCTG CTGCTCTACCCACTCGATGAA Wu et al. (2017)

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un-published of our group, the yield a common copio-trophic strain Escherichia coli BL21 growing with 13 mg C/L was 8 × 106 cells/mL (data not shown), which is 100 times lower than that of strain WZN-1.

Optimization of BDE209 degradation conditionsOrthogonal tests were designed to acquire optimal con-dition for biodegradation of BDE 209 (Additional file  1: Table S1). The results of the orthogonal tests are pre-sented in Additional file 1: Table S2. Range analysis and variance analysis of the orthogonal test were carried out to evaluate the influence of different factors on BDE 209 biodegradation (Additional file  1: Tables S2, S4). The optimum biodegradation conditions were pH 5, 25  °C, 0.5% salinity, 150 mL MSM volume. The results showed

that pH was the most important factor (R = 49.41), fol-lowed by MSM volume (R = 15.13), salinity (R = 13.49), temperature (R  =  10.33) and BDE 209 concentration (R = 6.51) (Additional file 1: Table S2). The variance anal-ysis demonstrated that the most significant factor for BDE 209 degradation was pH (p  =  0.01), followed by salinity (p  =  0.40), volume (p  =  0.44), temperature (p = 0.67) and BDE 209 concentration (p = 0.91) (Addi-tional file 1: Table S4).

Biodegradation of BDE 209 and kinetic analysisThe BDE 209 biodegradation by the strain WZN-1, as well as the bacterial growth were investigated simulta-neously in MSM under the optimal condition, to which 65  μg/L BDE 209 was applied as carbon and energy source. The results showed strain WZN-1 could biode-grade 55.15% of the initial BDE 209 in the batch experi-ments over a 30-d incubation period (Fig.  1), and the pH values remained at initial levels. The biomass of strain WZN-1 rapidly increased with time in the first 15 days incubation, and reached the maximum concen-tration of 2.50 ×  105 cells/mL, then decreased from 15 to 25 days and kept relatively stable from 25 to 30 days. Figure  2 shows the flow cytometric dot plots of strain WZN-1 growth measured during the biodegradation process. The results indicated that the cellular proper-ties (e.g. cell size and cellular nucleic acid content) of strain WZN-1 was consistent with the BDE 209 degra-dation rates. During the first 15 days, BDE 209 biodeg-radation efficiencies were higher concomitant with the rapid bacteria growth. After that, the strain gradually adapted to the BDE 209 environment, so that the BDE 209 degradation rate maintained a stable increase from 15 to 30 days, which means that the strain has a strong

Fig. 1 Biodegradation of BDE209 by strain WZN-1 under the optimized condition. The error bars represent standard deviation of triplicate measurements

0.1 1 10 100 10000.1

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Fig. 2 Examples of flow cytometric dot plots of strain WZN-1 at different time during BDE 209 degradation: a 0 day; b 15 days; c 30 days. FL1 repre-sents green fluorescence was collected at FL1 = 520 ± 20 nm, and FL3 represents the red fluorescence was collected at FL3 = 630 nm

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BDE 209 biodegradation capacity (Guangyu et al. 2013). The kinetic analysis indicated that the biodegradation of BDE 209 by strain WZN-1 well fitted the first-order kinetics (Additional file 1: Fig. S4). BDE 209 degradation rate constant was 0.0456 ±  0.0095/days, the coefficient of determination R2 was 0.7862, and the half-lives was 15.8869 ± 4.6807 days.

Identification of intermediate products and biodegradation pathwayTo identify the intermediate products and biodegrada-tion pathway of BDE 209 by strain WZN-1, the BDE 209 biodegradation products in different times were detected and identified by GC–MS (Additional file 1: Fig. S5). The proposed debromination pathway for BDE 209 biodeg-radation by WZN-1 strain is shown in Fig.  3. The GC–MS analysis confirmed that one nona-BDEs (BDE 208), one hepta-BDE (BDE 190), two hexa-BDEs (BDE 138 and BDE 154), one penta-BDE (BDE 85), two tetra-BDEs (BDE 47 and BDE 66), and one tri-BDE (BDE 28) were formed during BDE 209 degradation by strain WZN-1. This indicated that BDE 209 was first converted by para debromination to BDE 208 (Guangyu et al. 2013). Then, the BDE 208 was debrominated to BDE 190, and BDE 190 was converted by ortho and meta debromination to BDE 138 and BDE 154. After that, BDE 138 and BDE 154 were converted to BDE 85 by ortho and meta debromi-nation. BDE 66 and BDE 47 were formed by the ortho and meta debromination of BDE 85. BDE 66 was finally converted to BDE 28 by meta debromination, and BDE 47 converted to BDE 28 by ortho debromination. In com-parison, there was no detectable change of lower PBDEs congeners occurred over time in the control experiments (Additional file  1: Fig. S5). Furthermore, samples were full-scanned by UPLC–MS and the mass spectra were recorded in full scan mode (m/z 50–1000). The proposed OH-PBDEs and open-ring products based on UPLC–MS analysis were listed in Fig. 4. Four major biodegradation products were deducted, including three OH-PBDEs: OH-Octa-BDEs (m/z 817), OH-Hexa-BDEs (m/z 658), OH-Tri-PBDEs (m/z 423); and one possible ring opening product bromo-pentenal (m/z 163).

Genome analysis of strain WZN‑1 and real‑time qPCRAccording to the genome sequencing, the genome size of strain WZN-1 is 4,512,703  bp with a GC content of 66.62%. The genome contained 4137 genes, the total length of genes was 3,961,347 bp, which makes up 87.78% of genome (Wu et  al. 2017). Based on the functional annotation of the 4137 genes, four genes were identi-fied and supposed to have BDE 209 degradation abil-ity (Table  1). In this study, the real-time qPCR analysis showed that the expression of the four identified BDE 209 degradation genes was significantly induced and upregulated by BDE 209, demonstrating their involve-ment in BDE 209 degradation process (Fig.  5). The expression levels of the four genes were showed similar trend, where the level first increased and then decreased along the degradation process. This is consistent with the strain WZN-1 growth in BDE 209, which is rapidly increased in the first 5 days incubation, and reached the

Fig. 3 A proposed debromination pathway for BDE 209 biodegrada-tion by strain WZN-1

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maximum concentration of 1.60  ×  107 cells/mL, then decreased to 8.76 ×  106 and 2.31 ×  106 cells/mL on 10 and 20 days, respectively. Generally, the expression lev-els of dehalogenase gene 1 and dehalogenase gene 2 were significant higher than the hydrolase gene and dioxyge-nase gene. The dehalogenase gene 1 showed the highest relative expression (4.38-fold) on 5 days, decreased to 2.28-fold on 10 days, and back to the initial level on 20

days. However, the expression of the dehalogenase gene 2, hydrolase gene and dioxygenase gene were raised in the first 10 days, and showed the highest expression lev-els on 10 days, but their expressions were dropped on 20 days. The statistical analysis showed that all the three genes showed a significant expression level (p < 0.05) on 5 and 10 days. The expression of the four genes in strain WZN-1 enabled this strain to deplete BDE 209.

DiscussionPBDEs are persistent in various environments, and tend to accumulate in the biota (Mackintosh et  al. 2015). Despite the adverse environmental impact of decabro-modiphenyl ether (BDE 209), very limited information is available for its microbial degradation under aerobic conditions. The present study reported the aerobic bio-degradation of BDE 209 by a newly isolated strain, Sten-otrophomonas sp. strain WZN-1. In previous studies, Stenotrophomonas spp. has been found to be ubiquitous in marine, soil, rhizosphere of diverse plants (Denton and Kerr 1998) and polluted environments (Dungan et al. 2003). The genus Stenotrophomonas has been recorded as a promising candidate for biotechnological applica-tions involved in the detoxification of various man-made pollutants because its broad spectrum of metabolic prop-erties (Ryan et al. 2009), including utilization of polycy-clic aromatic hydrocarbons (Juhasz et al. 2000), benzene, toluene, ethylbenzene (Lee et al. 2002), chlorinated com-pounds (Somaraja et al. 2013), heavy metals (Ghosh and Saha 2013) and xenobiotics (Fuller et al. 2010). Therefore, Stenotrophomonas spp. has a great application value in the field of environmental remediation.

According to the orthogonal experiment results, pH value is a vital factor influencing the bacterial biodeg-radation ability of various organic pollutants. pH could adjust the metabolism and growth of the microorganisms via controlling the enzymatic activities (Jin et  al. 2012). The present study showed that strain WZN-1 preferred acidic environment at pH 5, where its BDE 209 degra-dation rate was significantly higher than those at other pH levels (Additional file 1: Table S2). Comparing to the effect of pH, the influence of temperature on the BDE 209 biodegradation by WZN-1 was not significant. It is consistent with a previous study, with the increasing tem-perature from 20 to 40 °C, the BDE 209 degradation rate ranged from 17.83 to 28.15%; the highest degradation rate was observed at 25  °C (Xin et  al. 2014). The MSM salinity affects the osmotic pressure, so that will affect the growth of bacteria and its degradation rate on BDE 209. Our results indicated that strain WZN-1 had consider-ate tolerance for NaCl over the range from 0 to 4%. The highest BDE 209 degradation rate of 29.47% occurred at 0.5% salinity. The volume of the MSM also apparently

Fig. 4 Possible intermediate products identified by UPLC–MS during BDE 209 biodegradation by strain WZN-1 under the optimal condi-tion: a OH-Octa-BDEs (m/z 817); b OH-Hexa-BDEs (m/z 658); c OH-Tri-BDEs (m/z 423) and bromo-pentanal (m/z 163)

Fig. 5 The relative expression of four BDE 209 degradation genes during the degradation process

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influences BDE 209 degradation rate by WZN-1 strain. With the volume increases from 50  mL to 250  mL, the BDE 209 degradation rate ranged from 16.14 to 31.18%, and the highest degradation rate was observed at 150 mL, which indicated that the oxygen plays a crucial role in the BDE 209 biodegradation system. Comparing to other fac-tors tested, the initial concentrations of BDE 209 had no significant impact (p = 0.91) on the BDE 209 biodegrada-tion efficiency by strain WZN-1.

Previously, researches on the biodegradation of BDE 209 and some other PBDEs congeners have been reported (Table 2). A number of bacteria including Pseu-domonas spp., Sphingomonas spp., Lysinibacillus fusi-formis DB-1, Bacillus cereus JP12, and Phlebia lindtneri JN45 have been demonstrated to transform and degrade different PBDE congeners (Kim et al. 2007; Lu et al. 2013; Xu and Wang 2014). A few microbial degradations of BDE 209 have also been reported. For instance, it was shown the Bacillus cereus JP12 could degrade over 88% of BDE 209 (initial concentration 1  mg/L) after 12 days incubation (Lu et al. 2013). Besides, 1 g/L bacterial cells of Enterococcus casseliflavus could degrade 52.2% of 1 mg/L BDE 209 in 7 days (Tang et al. 2016). The present study showed strain WZN-1 could biodegrade 55.15% of 65 μg/L BDE 209 over a 30-day incubation period under the optimized aerobic degradation conditions. Besides, strain WZN-1 exhibited high viability and was still cul-turable when exposed to 5 mg/L BDE 209 in MSM (data not shown). The biodegradation efficiency of BDE 209 by strain WZN-1 was comparable to the previous reports that microbial degradation of BDE 209 at μg/L to mg/L level (Table 2). The degradation ability of strain WZN-1 presented in this study could provide new options for the remediation of PBDEs.

Additionally, biodegradation of BDE 209 by strain WZN-1 was first-order kinetics. The half-life of BDE 209 was 15.8869  ±  4.6807 days, which revealed that BDE 209 could be rapidly degraded by strain WZN-1. PBDEs have very low water solubility and high octanol–water partition coefficient, hence they exhibited low biodegra-dation in different environment (Stiborova et  al. 2015). The microbial degradation kinetics of different PBDEs congeners in sediment and in sewage sludge have been investigated. For instance, Huang and colleagues found the anaerobic debromination rate constant for BDE 209 was 0.022/days, the half-life was 31.5 days in sediment (Huang et al. 2014). Another study also investigated the degradation of PBDEs in sediment (Yang et  al. 2015). Their results showed that the degradation rate con-stants and half-lives were 0.347/days and 2 days, 0.128/days and 5.4 days, 0.056/days and 12.4 days, 0.050/days and 13.9 days, 0.047/days and 14.8 days for BDE 15, BDE 28, BDE 47, BDE 99 and BDE 100, respectively.

Microbial degradation of BDE 209 in sewage sludge was much slower than that in sediment. It was reported that the degradation constant ranged between 0.00277 and 0.00379/days in sewage sludge under aerobic conditions (Stiborova et  al. 2015). So far, there are only few stud-ies described the biodegradation kinetics of BDE 209 in water. Comparing to previous studies, strain WZN-1 showed more effective biodegradation of BDE 209 in MSM with degradation rate constants 0.0456 ± 0.0095/days and half-life 15.8869 ± 4.6807 days.

As mentioned in previous studies, enzymatic dehalo-genation involves the removal of halogen atoms by hydrolytic, reductive, or oxygen dependent dehalogena-tion (Reineke 2001). Based on the observed lower PBDEs congeners, one of the major biodegradation mechanisms of BDE 209 by WZN-1 strain was debromination (Fig. 3, Additional file  1: Fig. S5). However, no bromide ions were detected by ion chromatography during BDE 209 degradation, which may due to the relatively low con-centrations of bromide released, or the concentration of bromide ions is lower than the detection limit of the instrument. To date, information on the aerobic degra-dation of BDE 209 was relatively limited. It was reported that deca-BDE could be aerobically debrominated to octa-BDE and hepta-BDE congeners (Deng et  al. 2011). In addition, lower brominated PBDEs such as BDE 196, BDE 197, BDE 202, BDE 203, and BDE 183 were detected during BDE 209 biodegradation (Guangyu et  al. 2013). The ring of fully brominated BDE 209 is difficult to open directly because of its symmetric molecular structure and enhanced chemical stability by higher bromine con-stitution. Therefore, the biodegradation of BDE 209 most likely started with debromination, and resulted in the generation of lower PBDEs congeners, which was then easier to be broken down via ring opening. In this study, lower PBDEs congeners were detected but only in very low amount. It may due to the intracellular enzymatic degradation of BDE 209 which happens in the bacterial cell (Tang et  al. 2016). Furthermore, other degradation mechanisms are likely to be involved in the biodegrada-tion process, such as the hydroxylation.

Hydroxylation reaction is an important step in the degradation of halo aromatic compounds by organisms (Wang et al. 2016). It was revealed that PBDEs could be transformed to hydroxylated metabolites in many organ-isms, such as fish, plants, rats, mice, and humans (Kim et  al. 2014; Shi et  al. 2015). In this study, the deducted OH-Octa-BDEs, OH-Hexa-PBDEs and OH-Tri-PBDEs suggested BDE 209 could be degraded into OH-PBDEs through hydroxylation. This finding is consistent with a former research that Burkholderia xenovorans LB400 convert most of a mono-BDE to a hydroxylated mono-BDE (Robrock et al. 2009). Besides, Cao et al. found BDE

Page 9 of 12Wu et al. AMB Expr (2018) 8:27

Tab

le 2

Co

mp

aris

on

of t

he

PB

DEs

deg

rad

atio

n e

ffici

ency

rep

ort

ed in

pre

vio

us

rep

ort

s

–: n

ot m

entio

n

PBD

EsTi

me

Init

ial c

once

ntra

tion

Deg

rad

atio

n b

acte

ria

Deg

rad

atio

n e

ffici

ency

Refe

ren

ces

BDE

209

30 d

ays

65 μ

g/L

Sten

otro

phom

onas

sp.

str

ain

WZN

-155

.15%

This

stu

dy

Dip

heny

l eth

er, 4

-bro

mo-

, 2,4

-dib

rom

o-,4

,4′-d

ibro

mod

iphe

nyl e

ther

8 da

ys1

g/L

diph

enyl

eth

erSp

hing

omon

as s

p. P

H-0

710

0% (d

iphe

nyl e

ther

), 23

% (4

-bro

mo)

, 14

% (2

,4-d

ibro

mop

heny

l eth

er),

8%

(4,4′d

ibro

mod

iphe

nyl e

ther

)

Kim

et a

l. (2

007)

Tetr

a-, p

enta

- and

hex

a-BD

Es3

days

17 n

g/m

LRh

odoc

occu

s jos

tii R

HA

1, B

urkh

olde

ria

xeno

vora

ns L

B400

, Pse

udon

ocar

dia

diox

aniv

oran

s CB1

190

RHA

1 tr

ansf

orm

ed g

reat

er th

an 9

0% o

f m

ono-

and

di-B

DE

cong

ener

s, LB

400

tran

sfor

med

10–

45%

of p

enta

-BD

E co

ngen

ers

Robr

ock

et a

l. (2

009)

Oct

a-BD

E2

mon

ths

45 n

M n

ona-

BDE,

181

nM

oct

a-BD

Es,

294

nM h

epta

-BD

E, a

nd 1

9 nM

hex

a-BD

E

Deh

aloc

occo

ides

spe

cies

–Le

e an

d H

e (2

010)

BDE

47, 9

9, 1

0014

day

s11

80 n

MD

ehal

ococ

coid

es a

nd D

esul

fovi

brio

spp

.88

–100

%Le

e et

al.

(201

1)

BDE

209

6 da

ys6

mg/

LLy

sinib

acill

us fu

sifor

mis

DB-

1<

20%

Den

g et

al.

(201

1)

BDE

209

90 d

ays

10 μ

M o

f BD

E 20

9Ps

eudo

mon

as s

pp.

< 1

2%Q

iu e

t al.

(201

2)

BDE

209

12 d

ays

1 m

g/L

Baci

llus c

ereu

s JP1

288

%Lu

et a

l. (2

013)

Oct

a an

d pe

nta-

BDE

tech

nica

l mix

ture

s21

wee

ks28

0 nM

PBD

Es in

an

octa

-BD

E m

ixtu

reAc

etob

acte

rium

sp.

str

ain

AG

Deh

alo-

cocc

oide

s sp.

str

ain

DG

96%

Din

g et

al.

(201

3)

BDE

472

wee

ks20

μg/

LPs

eudo

mon

as st

utze

ri97

.94%

Zhan

g et

al.

(201

3)

BDE

477

days

50 μ

g/L

Pseu

dom

onas

put

ida

sp. s

trai

n TZ

-149

.96%

Xin

et a

l. (2

014)

BDE

209

7 da

ys1

mg/

LEn

tero

cocc

us c

asse

liflav

us52

.2%

Tang

et a

l. (2

016)

BDE

209

12 d

ays

0.5

mg/

LSt

aphy

loco

ccus

hae

mol

ytic

us L

Y1 a

nd

Baci

llus p

umilu

s LY2

29.8

and

39.

2%W

ang

et a

l. (2

016)

BDE

209

BDE

209

144

h5

days

50 m

g/L

20 m

g/L

Rhod

ococ

cus s

p.Ps

eudo

mon

as a

erug

inos

a65

.1%

85.1

2%Li

u et

al.

( 201

5, 2

016)

Page 10 of 12Wu et al. AMB Expr (2018) 8:27

47 was more easily to transform via hydroxylation by Phanerochaete chrysosporium (Cao et al. 2017). The m/z 163 was proposed to be bromo-pentanal indicated ring opening process may occurred in the BDE 209 aerobic degradation by strain WZN-1. Previously, Xu and Wang found Phlebia lindtneri JN45 could degrade BDE 209 through debromination, hydroxylation, and ring-opening reactions (Xu and Wang 2014). BDE 209 degradation by Bacillus pumilus LY2 were supposed to be three possi-ble metabolic pathways: debromination, hydroxylation, and cleavage of the diphenyl ether bond (Wang et  al. 2016). BDE 209 has a symmetric molecular structure and good chemical stability. Therefore, the biodegradation of BDE 209 most likely started with debromination and hydroxylation, and resulted in the generation of lower PBDEs congeners or OH-PBDEs, which was then easier to be broken down via ring opening. The present study indicated that BDE 209 was successively biodegraded to BDE 28 via a debromination process, and the proposed OH-PBDEs and bromo-pentanal supported the possible of hydroxylation and ring opening process. The results indicated further mineralization of in the tricarboxylic acid cycle (TCA) could be possible. Further researches of BDE 209 degradation by strain WZN-1 requires using standard substances verify these possible hydroxylation and ring opening products.

So far, very little is known about aerobic microbial transformation of PBDEs by microorganisms and the enzymes involved in PBDEs transformation (Robrock et al. 2011). In the present study, the significant upregu-lation of the four BDE 209 degradation genes confirmed the debromination, hydroxylation and ring opening process of BDE 209 degradation by strain WZN-1. The functional annotation of the genome showed that the hydrolase gene is an alpha/beta hydrolase in NR data-base. The KEGG database suggested this gene is dhaA haloalkane dehalogenase, and has xenobiotics biodegra-dation and metabolism function. The dioxygenase gene in strain WZN-1 was identified as an aromatic ring-open-ing dioxygenase via COG database, which suggested the potential ring opening ability of strain WZN-1. Moreo-ver, the dehalogenase gene 1 and dehalogenase gene 2 were identified as haloacid dehalogenases of Stenotropho-monas maltophilia from the NR database. Previous study showed that enzymes were involved in the PCB degrada-tion, transport processes, energetic metabolism, electron transport, and carbon metabolism(Zhang et  al. 2014). Dioxygenase enzymes are responsible for degradation of diphenyl ether and may be responsible for degradation of less brominated PBDEs (Lee et al. 2011). Additionally, the 2D electrophoresis (2-DE) and matrix-assisted laser desorption/ionization time of flight mass spectrometry (MALDI-TOF MS) were used to identify upregulated

proteins during PCB degradation, genes encoding for dioxygenase, ABC transporters, transmembrane pro-teins, electron transporter, and energetic metabolism proteins were significantly upregulated. Similarly, Tang et  al. also suggested that enzymes and genes might act as functional protein in BDE 209 biodegradation, ATP synthase and ABC transporter permease played a criti-cal role in BDE 209 transport from extracellular environ-ment to intracellular cells (Tang et al. 2016). Besides the four identified BDE 209 degradation genes, we also found twelve of ATP synthases, thirty-six of ABC transporter permeases, nineteen of hydroxylases, and sixty-five of oxygenase in strain WZN-1 (data not shown), which indicates that there may more genes involved in the bio-degradation of BDE 209 by strain WZN-1.

In summary, a newly isolated bacterium Stenotropho-monas sp. strain WZN-1 could degrade 55.15% of 65  μg/L BDE 209 in 30 days under the optimized con-dition (i.e. pH 7, 25 °C, 0.5% salinity, 150 mL MSM vol-ume). The degradation kinetic was fitted to a first-order kinetics model. The identified lower PBDEs congeners, OH-PBDEs and bromo-pentanal of the degradation of BDE 209 by WZN-1 suggested the degradation pathways included the debromination, and proposed hydroxyla-tion and ring opening processes. Four identified genes was significantly induced and up-regulated by BDE 209, which indicated that these genes are of great importance in the BDE 209 degradation process. These findings could provide useful information for PBDE remediation appli-cations and contribute to a better understanding of BDE 209 biodegradation process.

AbbreviationsBDE 209: decabromodiphenyl ether; PBDEs: polybrominated diphenyl ethers; POPs: persistent organic pollutants; THF: tetrahydrofuran; MSM: minimal salt medium; BMM: basal mineral medium; LB: Luria–Bertani; SEM: scanning elec-tron microscopy; FCM: flow cytometry; NCI: negative chemical ionization; ESI: negative electrospray ionization mode.

Authors’ contributionsZW conducted research work and drafted manuscript. MX assisted in conduct-ing experiments. YL participated in experimental design and data analysis. GG carried out the statistical design and data analysis. MB contributed to the data analysis and manuscript editing. YW designed and supervised all experiments and review of manuscript. All authors read and approved the final manuscript.

Author details1 Key Laboratory of Pollution Processes and Environmental Criteria (Ministry of Education), Tianjin Key Laboratory of Environmental Remediation and Pol-lution Control, College of Environmental Science and Engineering, Nankai University, Tianjin 300350, China. 2 State Key Laboratory of Medicinal Chemical Biology, Nankai University, Tianjin 300350, China. 3 College of Life Sciences, Nankai University, Tianjin 300071, China.

Additional file

Additional file 1. Additional figures and tables.

Page 11 of 12Wu et al. AMB Expr (2018) 8:27

Competing interestsThe authors declare that they have no competing interests.

Ethics approval and consent to participateNot applicable.

FundingThis work was supported by National Key Basic Research Development Program (2015CB459000), National Natural Science Foundation of China (31670498, 31322012), the Tianjin Municipal Science and Technology Commis-sion (13JCYBJC20800) and the Fundamental Research Funds for the Central Universities.

Publisher’s NoteSpringer Nature remains neutral with regard to jurisdictional claims in pub-lished maps and institutional affiliations.

Received: 24 January 2018 Accepted: 16 February 2018

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